Pharmaceuticals and Personal Care Products (Ppcps): a Review on Environmental Contamination in China

Environment International 59 (2013) 208–224

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Review Pharmaceuticals and personal care products (PPCPs): A review on environmental contamination in China

Jin-Lin Liu b, Ming-Hung Wong a,b,⁎ a School of Environmental and Resource Sciences, Zhejiang Agricultural and Forestry University, Linan, PR China b Croucher Institute for Environmental Sciences, and Department of Biology, Hong Kong Baptist University, Hong Kong, PR China article info abstract

Article history: Pharmaceuticals and personal care products (PPCPs) which contain diverse organic groups, such as antibi- Received 15 February 2013 otics, hormones, antimicrobial agents, synthetic musks, etc., have raised signiﬁcant concerns in recently Accepted 18 June 2013 years for their persistent input and potential threat to ecological environment and human health. China is Available online 6 July 2013 a large country with high production and consumption of PPCPs for its economic development and population growth in recent years. This may result in PPCP contamination in different environmental media of Keywords: China. This review summarizes the current contamination status of different environment media, including Pharmaceuticals and personal care products (PPCPs) sewage, surface water, sludge, sediments, soil, and wild animals, in China by PPCPs. The human body burden China and adverse effects derived from PPCPs are also evaluated. Based on this review, it has been concluded that Contamination more contamination information of aquatic environment and wildlife as well as human body burden of PPCPs Sewage in different areas of China is urgent. Studies about their environmental behavior and control technologies Surface water need to be conducted, and acute and chronic toxicities of different PPCP groups should be investigated for assessing their potential ecological and health risks. © 2013 Elsevier Ltd. All rights reserved.

Contents

1. Introduction ...... 209 1.1. Classification of PPCPs ...... 209 1.2. Sources, fates, and persistence of PPCPs ...... 210 2. PPCP production and usage in China ...... 210 3. Occurrence of PPCPs in different environmental media of China ...... 210 3.1. PPCPs in sewage and sludge ...... 210 3.1.1. PPCPs in sewage ...... 210 3.1.2. Fates of PPCPs in STPs ...... 214 3.1.3. Removal efficiencies of PPCPs in STPs ...... 214 3.1.4. PPCPs in sludge ...... 214 3.2. PPCPs in surface water and sediments ...... 214 3.2.1. PPCPs in surface water ...... 214 3.2.2. Fates of PPCPs in surface water ...... 217 3.2.3. PPCPs in sediments ...... 217 3.2.4. Removal efficiencies of PPCPs in water treatment plants ...... 217 3.3. PPCPs in soil ...... 218 3.4. PPCPs in wild animals ...... 218 4. Human exposure and body loading of PPCPs ...... 218 4.1. Human exposure ...... 218 4.2. Body loading ...... 218 5. Toxicity and risk assessments of PPCPs ...... 219 5.1. Toxicity ...... 219 5.2. Risk assessments ...... 220

⁎ Corresponding author at: School of Environment and Energy, Peking University, Shenzhen, PR China. Tel.: +852 3411 7746; fax: +852 3411 7743. E-mail address: [email protected] (M.-H. Wong).

6. Regulations and control strategies of PPCP contamination ...... 220 7. Conclusions ...... 221 Acknowledgments ...... 221 References ...... 221

1. Introduction lipid regulators, β-blockers, contrast media, and cytostatic drugs for pharmaceuticals; and antimicrobial agents, synthetic musks, insect Pharmaceuticals and personal care products (PPCPs) have received repellants, preservatives, and sunscreen UV filters for personal care growing attention in recent years as emerging contaminants for their products (Table 1)(Daughton and Ternes, 1999). Among the pharma- possible threats to aquatic environment and human health. They con- ceutical group, antibiotics have received special attention for their tain a large and diverse group of organic compounds, including pharma- wide application in human therapy and livestock agriculture. Persistent ceutical drugs and integrates of daily personal care products (PCPs), exposure of antibiotics can result in the emergence of resistant bacteria such as soaps, lotions, toothpaste, fragrances, sunscreens, etc., which strains with public health concerns (Zhang et al., 2009b). Antibiotics are widely used in high quantities throughout the world, together contain several subgroups, such as macrolides (e.g. erythromycin, with their metabolites and transformation products (Daughton and roxithromycin), sulfonamides (e.g. sulfamethoxazole, sulfadimethoxine), Ternes, 1999; Kummerer, 2000). As an important group of organic pol- and fluoroquinolones (e.g. norfloxacin, ciprofloxacin) (Heberer, 2002). lutants with intensive studies in recent years, PPCPs have been found Hormones are another most studied group of pharmaceuticals which to be ubiquitous in the aquatic environment throughout the world are believed to be connected with the endocrine disrupting effects of (Carballa et al., 2004; Kasprzyk-Hordern et al., 2009; Lishman et al., polluted water bodies (Lai et al., 2002). The most concerned and studied 2006; Nakada et al., 2006; Tauxe-Wuersch et al., 2005), including in hormones are steroid estrogens, including natural steroid estrogens the raw water sources of drinking water treatment plants (Radjenovic which are primarily excreted by humans and animals, e.g. estrone (E1), et al., 2008; Ternes et al., 2002; Vieno et al., 2007a). estradiol (E2), estriol (E3), and synthetic steroid estrogens which are China is a large country with high production and consumption of used as oral contraceptives, mainly ethinylestradiol (EE2) (Heberer, PPCPs owe to its economic development and population growth, 2002). Natural steroid estrogens are not actually pharmaceuticals, how- which may result in serious pollution by PPCPs. This review aims to ever, they are usually studied together with synthetic hormones for summarize the current PPCP contamination status in different envi- their endocrine disrupting effects in polluted water. They will be also ronmental media of China and assess their possible threats to ecosys- discussed in the current review. Other pharmaceutical groups include tem and human health. analgesics and anti-inflammatory drugs (such as diclofenac and ibuprofen); antiepileptic drugs (such as carbamazepine and primidone); 1.1. Classification of PPCPs blood lipid regulators (such as clofibrate and gemfibrozil); β-blockers (such as metoprolol and propanolol); and contrast media (such as PPCPs contain diverse groups of organic compounds, such as anti- iopromide and diatrizoate) (Heberer, 2002). For the groups of personal biotics, hormones, anti-inflammatory drugs, antiepileptic drugs, blood care products, triclosan and triclocarban are the two typical antimicrobial

Table 1 Classiﬁcation of PPCPs.

Subgroups Representative compounds

Pharmaceuticals Antibiotics Clarithromycin Erythromycin Sulfamethoxazole Sulfadimethoxine Ciprofloxacin Norfloxacin Chloramphenicol Hormones Estrone (E1) Estradiol (E2) Ethinylestradiol (EE2) Analgesics and anti-inflammatory drugs Diclofenac Ibuprofen Acetaminophen Acetylsalicylic acid Antiepileptic drugs Carbamazepine Primidone Blood lipid regulators Clofibrate Gemfibrozil β-blockers Metoprolol Propanolol Contrast media Diatrizoate Iopromide Cytostatic drugs Ifosfamide Cyclophosphamide Personal Care Products Antimicrobial agents/Disinfectants Triclosan Triclocarban Synthetic musks/Fragrances Galaxolide (HHCB) Toxalide (AHTN) Insect repellants N,N-diethyl-m-toluamide (DEET) Preservatives Parabens (alkyl-p-hydroxybenzoates) Sunscreen UV filters 2-ethyl-hexyl-4-trimethoxycinnamate (EHMC) 4-methyl-benzilidine-camphor (4MBC) 210 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 agents frequently detected in wastewater. The synthetic musks 3.0 include nitro musks (mainly musk xylene (MX) and musk ketone Pharmaceuticals (a) (MK)) and polycyclic musks (mainly galaxolide (HHCB) and toxalide 2.5 (AHTN)) with more production and application than the nitro group in recent years. N,N-diethyl-m-toluamide (DEET) is the main active 2.0 ingredient of insect repellents and regularly detected. Parabens are typical preservatives, and 2-ethyl-hexyl-4-trimethoxycinnamate 1.5 (EHMC) and 4-methyl-benzilidine-camphor (4MBC) are sunscreen UV fi 1.0

lter species (Brausch and Rand, 2011). (million tonnes)

1.2. Sources, fates, and persistence of PPCPs 0.5

0.0 The main source of PPCP infusion into the environment is through 2003 2004 2005 2006 2007 2008 2009 2010 1-11 2011 sewage treatment plants (STPs) (Daughton and Ternes, 1999). The Active pharmaceutical ingredient production presence of PPCPs in wastewater treatment plants had been reported in different countries all over the world, mostly in the levels of ng L−1 25 to μgL−1, such as USA (Boyd et al., 2004), United Kingdom (Ashton et Personal Care Products (b) al., 2004), Spain (Carballa et al., 2004), Finland (Lindqvist et al., 2005), 20 and Japan (Nakada et al., 2006). Studies also reveal that the removal efficiency of PPCPs by the conventional wastewater treatment processes (flocculation, sedimentation, and active sludge treatment) is 15 limited (Castiglioni et al., 2006; Lishman et al., 2006; Paxeus, 2004; Santos et al., 2007). 10 After discharging from STPs, PPCPs in sewage would cause subse- quent contamination to the receiving water bodies. For their polar 5 properties, the fraction which enters into the atmospheric environ- Market scale (billion USD) ment will be limited. Their distribution will primarily occur in the 0 aquatic environment. PPCPs may also be adsorbed onto the active 2007 2008 2009 2010 2011 (Estimated) 2012 (Estimated) 2013 sludge in STPs and then introduced into the environment through sludge land application (Daughton and Ternes, 1999). As a group of novel emerging contaminants, PPCPs show varied properties with the conventional persistent organic pollutants whose sources may have been banned or limited. The input of PPCPs with STPs as the Fig. 1. (a) Active pharmaceutical ingredient production of China (2003–2011) (SERI, – main source is perpetual, resulting in a steady-state concentration in 2012). (b) Personal care product market scale of China (2007 2013) (BosiData, 2011). aquatic systems, which has been described as “pseudo-persistent” in human therapy have been applied in livestock agriculture (Heddini (Daughton, 2003). Persistent exposure by PPCPs even at low concentra- et al., 2009; Li and Wang, 2009; Zheng and Zhou, 2007). The emergence tion levels can be significant. or persistence of resistant bacteria strains derived from antibiotic abuse may become a serious public health crisis in the future. 2. PPCP production and usage in China China is also among the top three countries with the largest personal care product consumption, together with America and Japan (ChinaIRN, China has become a large country with high production and usage 2012). China possesses the fastest growing rate of personal care product volumes of pharmaceuticals as well as a rapid growing rate of personal market in the world, which will reach 8% between 2010 and 2013. In care product consumption, which may result in significant occurrence 2013, the total industry value will reach USD21.3 billion, accounting of PPCPs in the environment. for about 10% of the world (Fig. 1(b)) (BosiData, 2011; ChinaIRN, 2012). The production of active pharmaceutical ingredients by China has rapidly increased in recent years, and in fact China has become the largest producer of active pharmaceutical ingredients of the world. As 3. Occurrence of PPCPs in different environmental media of China shown in Fig. 1(a), from 2003 to 2011, the production of active pharmaceutical ingredients by China has increased two times, and in 2011 3.1. PPCPs in sewage and sludge about two million tonnes of pharmaceuticals were produced. The pharmaceutical production by China can account for more than 20% of the Sewage treatment plants are considered as the main source of total production volume of the world (SERI, 2012). More than 1500 PPCPs. Occurrence of PPCPs in STPs receives great attention, from kinds of active pharmaceutical ingredients were produced, and more which the information about environmental loads of PPCPs can be than 6900 pharmaceutical-manufacturing companies were registered obtained. Their fate in STPs also matters. Whether or not PPCPs can in 2007 (Zhou et al., 2010). China also has the highest production of sev- be removed by the wastewater treatment processes determines eral antibiotic species, such as penicillin and terramycin (SCIO, 2008). their final input into the receiving aquatic environment. A number Besides the huge pharmaceutical production volume, the consump- of studies in China focused on the occurrence and fate of PPCPs in tion rate is also remarkable, especially for the severe antibiotic abuse in STPs, including sewage (Table 2) and sludge. Most investigations current China. The average usage of antibiotics by Chinese is 10 times were carried out in several hotspot study areas, such as the Pearl more than the usage by Americans (CAST, 2008). Around 75% of the pa- River Delta, Yangtze River Delta, and Beijing area, due to their dense tients and 80% of the inpatients with seasonal influenza are prescribed population and rapid socio-economic development. antibiotics (Heddini et al., 2009; Zheng and Zhou, 2007). For the treatment of animal diseases and promotion of livestock growth, antibiotics 3.1.1. PPCPs in sewage are also widely used in livestock agriculture. The antibiotic usage volume in livestock farms is about 97,000 tonnes and can account for 3.1.1.1. Antibiotics. As the most concerned group of pharmaceuticals, 46% of the total volume every year, and most antibiotic species used antibiotics had been found ubiquitously in STP sewage of China, J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 211

Table 2 Occurrence of PPCPs in sewage of STPs.

Location Chemical Concentration Media Reference

Guangzhou Antibiotics 1730–7910 ng L−1 Sewage influent Peng et al. (2006) ND−9460 ng L−1 Sewage effluent Guangzhou Antibiotics b100–5597 ng L−1 Sewage influent Peng et al. (2008a) b35–1920 ng L−1 Sewage effluent Guangzhou Antifungal drugs 1–1834 ng L−1 Sewage influent and effluent Peng et al. (2012) Guangzhou and Hong Kong Antibiotics 10–1978 ng L−1 Sewage influent Xu et al. (2007a) 9–2054 ng L−1 Sewage effluent Hong Kong Antibiotics ND−7870 ng L−1 Sewage effluent Minh et al. (2009) Hong Kong Antibiotics 3.2–1718 ng L−1 Sewage influent Li and Zhang (2011) 1.3–1176 ng L−1 Sewage effluent Hong Kong Antibiotics b4–2900 ng L−1 Sewage influent Gulkowska et al. (2008) b4–1800 ng L−1 Sewage effluent Shenzhen Antibiotics 29–1100 ng L−1 Sewage influent Gulkowska et al. (2008) Jiangsu Antibiotics Max. 211000 ng L−1 Livestock sewage Wei et al. (2011) Hangzhou Antibiotics 108–1405 ng L−1 Sewage influent Tong et al. (2011) 54–429 ng L−1 Sewage effluent Chongqing Antibiotics ND−4240 ng L−1 Hospital sewage Chang et al. (2010) 7–157 ng L−1 Swine nursery sewage 11–825 ng L−1 Slaughter house sewage 18–2020 ng L−1 Municipal sewage Beijing Antibiotics Max. 3100 ng L−1 Sewage influent Gao et al. (2012) Max. 1200 ng L−1 Sewage effluent Beijing Antibiotics ND−2900 ng L−1 Sewage influent Chang et al. (2008) ND−4000 ng L−1 Sewage effluent Beijing Antibiotics 28–1208 ng L−1 Sewage influent Xiao et al. (2008) 6–503 ng L−1 Sewage effluent Beijing Antibiotics 600–32700 ng L−1 Swine sewage Ben et al. (2008) Guangzhou Hormones 1.2–1368 ng L−1 Sewage influent Liu et al. (2012b) 1.0–23.1 ng L−1 Sewage effluent Guangzhou Hormones 2.4–621 ng L−1 Sewage influent Liu et al. (2011b) b0.4–77.3 ng L−1 Sewage effluent Wuhan Hormones 4.8–82.4 ng L−1 Sewage influent Jin et al. (2008) ND-31.5 ng L−1 Sewage effluent Beijing Hormones 73–155 ng L−1 Sewage of a controceptives producing factory Cui et al. (2006) Beijing Hormones ND−4100 ng L−1 Sewage influent and effluent Zhou et al. (2011e) Beijing Hormones ND−760 ng L−1 Sewage influent Zhou et al. (2011d) ND−74.2 ng L−1 Sewage effluent Beijing Hormones 22.4–911.5 ng L−1 Sewage influent Zhou et al. (2009b) ND−253.8 ng L−1 Sewage effluent Beijing Hormones ND−3700 ng L−1 Sewage influent Chang et al. (2011) ND−12 ng L−1 Sewage effluent Beijing Hormones b0.08–120 ng L−1 Sewage influent Chang et al. (2007) b0.02–1.9 ng L−1 Sewage effluent Tianjin Hormones b0.4–8.1 ng L−1 Sewage influent Wang et al. (2005) b0.4–3.4 ng L−1 Sewage effluent Guangzhou Pharmaceuticals b0.7–3239 ng L−1 Sewage effluent Zhao et al. (2010b) Shanghai Carbamazepine 230–1110 ng L−1 Sewage influent and effluent Zhou et al. (2011c) Beijing Pharmaceuticals 38–89500 ng L−1 Sewage influent Zhou et al. (2010) 27–2700 ng L−1 Sewage effluent Beijing Pharmaceuticals 4.4–6600 ng L−1 Sewage influent Sui et al. (2010) 2.2–320 ng L−1 Sewage effluent Beijing Pharmaceuticals ND−11400 ng L−1 Sewage influent Sui et al. (2011) ND−1020 ng L−1 Sewage effluent Taiwan Pharmaceuticals ND−30967 ng L−1 Sewage influent Lin et al. (2010) Guangzhou Antimicrobial agents 10.9–342 ng L−1 Sewage effluent Zhao et al. (2010a) Guangzhou Synthetic musks ND−3190 ng L−1 Sewage influent Zeng et al. (2007) b10–2050 ng L−1 Sewage effluent Guangzhou Synthetic musks 3730–595480 ng L−1 Sewage influent Chen et al. (2007) 500–33540 ng L−1 Sewage effluent of a cosmetic plant Shanghai Synthetic musks b4–3430 ng L−1 Sewage influent Zhang et al. (2008) b4–336 ng L−1 Sewage effluent Beijing Synthetic musks 111.9–3003.8 ng L−1 Sewage influent Zhou et al. (2009a) 47.3–1258.3 ng L−1 Sewage effluent Beijing Insect repellents 124–1520 ng L−1 Sewage influent Sui et al. (2011) 21.6–235 ng L−1 Sewage effluent Tianjin UV filters 34–2128 ng L−1 Sewage influent Li et al. (2007) 21–1287 ng L−1 Sewage effluent

Other counties Korea Antibiotics ND−19401 ng L−1 Sewage influent Behera et al. (2011) (Ulsan) ND−21278 ng L−1 Sewage effluent U.K. Antibiotics b3–10025 ng L−1 Sewage influent Kasprzyk-Hordern et al. (2009) (South Wales) b3–3052 ng L−1 Sewage effluent Antibiotics Lindberg et al. (2005)

(continued on next page) 212 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224

Table 2 (continued) Location Chemical Concentration Media Reference

Sweden ND−2480 ng L−1 Sewage influent (5 cities) ND−1340 ng L−1 Sewage effluent Other counties Finland Antibiotics ND−4230 ng L−1 Sewage influent Vieno et al. (2007b) (12 cities) ND−130 ng L−1 Sewage effluent U.S. Antibiotics ND−1400 ng L−1 Sewage influent Brown et al. (2006) (New Mexico) 110–310 ng L−1 Sewage effluent U.S. Antibiotics 220–1140 ng L−1 Sewage influent Yang and Carlson (2003) (Colorado) 90–320 ng L−1 Sewage effluent Korea Hormones ND−802 ng L−1 Sewage influent Behera et al. (2011) (Ulsan) ND−24 ng L−1 Sewage effluent Japan Hormones 13.3–255 ng L−1 Sewage influent Nakada et al. (2006) (Tokyo) 0.3–110 ng L−1 Sewage effluent Portugal Hormones 103–2484 ng L−1 Sewage influent Salgado et al. (2010) (5 cities) ND−25 ng L−1 Sewage effluent Canada Hormones 2.4–78 ng L−1 Sewage influent Servos et al. (2005) (18 cities) 0.2–96 ng L−1 Sewage effluent Korea Pharmaceuticals ND−11239 ng L−1 Sewage influent Behera et al. (2011) (Ulsan) ND−5911 ng L−1 Sewage effluent Japan Pharmaceuticals 1.6–19400 ng L−1 Sewage influent Nakada et al. (2006) (Tokyo) 0.2–968 ng L−1 Sewage effluent U.K. Pharmaceuticals ND−482687 ng L−1 Sewage influent Kasprzyk-Hordern et al. (2009) (South Wales) ND−97616 ng L−1 Sewage effluent Portugal Pharmaceuticals 32–106490 ng L−1 Sewage influent Salgado et al. (2010) (5 cities) 20–43653 ng L−1 Sewage effluent Finland Pharmaceuticals 30–1350 ng L−1 Sewage influent Vieno et al. (2006) (3 cities) b5.8–1070 ng L−1 Sewage effluent Norway Pharmaceuticals ND−104000 ng L−1 Sewage influent Weigel et al. (2004) (Tromsø) ND−126000 ng L−1 Sewage effluent U.S. Pharmaceuticals 330–43800 ng L−1 Sewage influent Thomas and Foster (2005) (3 cities) ND−72 ng L−1 Sewage effluent Japan Synthetic musks 280–1400 ng L−1 Sewage influent Nakata and Shinohara (2010) (A town in western Japan) 27–660 ng L−1 Sewage effluent UV filters b8.7–78 ng L−1 Sewage influent 2.1–b 8.7 ng L−1 Sewage effluent U.K. Antimicrobial agents 27–65381 ng L−1 Sewage influent Kasprzyk-Hordern et al. (2009) (South Wales) 13–888 ng L−1 Sewage effluent Preservatives b2–30688 ng L−1 Sewage influent b0.6–155 ng L−1 Sewage effluent UV filters 9–13248 ng L−1 Sewage influent b2–6325 ng L−1 Sewage effluent Portugal Synthetic musks 1–11463 ng L−1 Sewage influent Salgado et al. (2010) (5 cities) 1–889 ng L−1 Sewage effluent Norway Triclosan 380–430 ng L−1 Sewage influent Weigel et al. (2004) (Tromsø) 160–480 ng L−1 Sewage effluent U.S. Synthetic musks 304–12700 ng L−1 Sewage influent Reiner et al. (2007a) (New York State) 495–3730 ng L−1 Sewage effluent ⁎ ND: not detected. with concentration up to several μgL−1 for most reviewed cases (ND As another developed area located in eastern China, antibiotics (not detected)-7910 ng L−1 for influent samples and ND-9460 ng L−1 were also present in STPs of Yangtze River Delta. Veterinary antibi- for effluent samples). In Pearl River Delta cities located in South China otics, including oxytetracycline, tetracycline, sulfamethazine, sulfadi- (such as Guangzhou, Shenzhen, and Hong Kong), antibiotic species azine, and sulfamethoxazole, were monitored in wastewater of 27 with high detection frequency and concentration included macrolides animal farms from Jiangsu Province, with the highest concentration −1 (e.g. roxithromycin and erythromycin–H2O(themaindegradationprod- up to 211 μgL (Wei et al., 2011). This is much higher than uct of erythromycin)), fluoroquinolones (e.g. ofloxacin and norfloxacin), the level in municipal wastewater, such as the maximum of − and sulfonamides (e.g. sulfamethoxazole) (Gulkowska et al., 2008; Xu et 1405 ng L 1 for sewage influent samples in STPs of Hangzhou locat- al., 2007a). Azole antifungal drugs, such as fluconazole, clotrimazole, ed in the same region (Tong et al., 2011),implyingtheimportanceof econazole, ketoconazole, and miconazole, also occurred in STPs (Huang livestock wastewater as the source of antibiotic input into the et al., 2012; Peng et al., 2012). Different cities have different antibiotic environment. composition patterns for sewage influents, e.g. cephalexin was the dom- Groups of antibiotics also existed in STPs of Beijing, the capital city inant species in Hong Kong sewage, while cefotaxim prevailed in of the nation located in North China. Veterinary antibiotics were − Shenzhen STPs (Gulkowska et al., 2008). This can be attributed to the re- found with a concentration range of 0.6–32.7 μgL 1 in swine waste- gional difference in antibiotic prescription patterns. The environmental water (Ben et al., 2008). This is the highest antibiotic level found in loads of antibiotics can be obtained according to the occurrence investi- the sewage of Beijing, exceeding the levels generally up to several gation results. According to the results from four sewage treatment μgL−1 for other cases (Chang et al., 2008; Gao et al., 2012; Xiao et plants in Guangzhou and Hong Kong, environmental loads of different al., 2008). This confirms the heavy loads of antibiotics from livestock antibiotic species were estimated to range from 0.5 to 828 g per day farms. (Xu et al., 2007a). From seven sewage treatment plants of Hong Kong, Generally, the sewage antibiotic levels of China are comparable the total input of antibiotics into the receiving Victoria Harbor was esti- with other areas of the world, such as other Asian countries (Behera matedtobe14.4kgperday(Minh et al., 2009). et al., 2011), Europe (Kasprzyk-Hordern et al., 2009; Lindberg et al., J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 213

2005; Vieno et al., 2007b), and U.S. (Brown et al., 2006; Yang and in wastewater. Besides the mostly studied estrogens, other hormone Carlson, 2003), up to several μgL−1. However, there is some differ- groups, such as androgens and progestagens, may have much higher ence between specific cases. Sewage antibiotic concentrations more contents as indicated by several studies (Chang et al., 2011; Liu et than 10,000 ng L−1 were found in Korea (Behera et al., 2011)andU.K. al., 2012b). This can be explained by their higher excretion amounts (Kasprzyk-Hordern et al., 2009) which are relatively higher than that in human urine than those of estrogens (Shemesh and Shore, 2003) of most studies in China, while sewage from Guangzhou and Hong and wide application in human and veterinary therapies, especially Kong (Minh et al., 2009; Peng et al., 2006) appeared to have more synthetic progestogens (Chang et al., 2011). Their high occurrence antibiotics (ND-9460 ng L−1) than the levels (ND-4230 ng L−1)from in sewage, sometimes more than several μgL−1, may pose potential Sweden (Lindberg et al., 2005), Finland (Vieno et al., 2007b), and U.S. harm to the ecosystem. However, due to the analytical difficulty, (Brown et al., 2006; Yang and Carlson, 2003). In China, the predominant studies about these hormone groups are relatively lacking. and frequently detected antibiotic groups, such as macrolides, fluoroquinolones, and sulfonamides, are related to the therapy use pattern 3.1.1.3. Other pharmaceuticals. Besides antibiotics and hormones, other and probably area-specific. Considering the intensive usage of veteri- pharmaceuticals were also present in STP sewage of China. Pharmaceu- nary antibiotics in livestock industry (Heddini et al., 2009; Li and tical groups, including anti-inflammatory drugs, antiepileptic drugs, Wang, 2009; Zheng and Zhou, 2007) and their significant occurrence blood lipid regulators, anti-hypertensive drugs, anticonvulsants, stimu- in livestock wastewater as indicated by several studies (Ben et al., lants, and antipsychotic drugs, had concentrations up to 89,500 ng L−1 2008; Wei et al., 2011), loads of veterinary antibiotics from livestock in influents and 3239 ng L−1 in effluents (Sui et al., 2010, 2011; Zhao et agriculture may pose a risk to the receiving water environment. al., 2010b; Zhou et al., 2010). The species with high detection frequency and concentration include caffeine, ibuprofen, diclofenac, carbamaze- 3.1.1.2. Hormones. Hormones as another important group of organic pine, etc. for their wide application, especially caffeine which can act compounds with public health concerns were also found in STP sewage as a tracer of domestic sewage pollution for its widespread occurrence of China. In most cases their concentrations were below the μgL−1 (Siegener and Chen, 2002). grade, and therefore relatively lower than that of antibiotics. Steroid es- Besides Mainland, Taiwanese consumers also use large quantities trogens, including the natural E1, E2, E3 and synthetic EE2, were mostly of pharmaceuticals, and 3.6 tonnes of them are estimated to be studied (Jin et al., 2008; Zhou et al., 2011d, 2011e). Other novel groups dumped annually (Lin et al., 2010). The study about pharmaceuticals of hormones, such as androgens, progestagens, and glucocorticoids, in STP sewage of Taipei, Taichung, and Tainan indicated cephalexin, were also investigated (Chang et al., 2011; Liu et al., 2012b). caffeine, acetaminophen, ibuprofen, and naproxen had the highest in- In Pearl River Delta, steroid estrogen species, such as estrone and fluent contents, ranging from 458 to 30,967 ng L−1; and ibuprofen 17α-estradiol, had occurrence range of 5.0–66.8 ng L−1 in sewage and naproxen had the highest effluent contents, ranging from 20 to samples of Guangzhou STPs (Yu et al., 2011b). The investigation 3807 ng L−1 (Lin et al., 2010). about steroid estrogens, androgens, progestagens, and glucocorti- These studies imply the existence of diverse pharmaceutical coids in STPs of Guangdong Province implied the androgen group groups in STP sewage of China. Levels of some species with high had the highest total concentration (13.3–1778 ng L−1), followed usage quantities, such as caffeine, may exceed that of antibiotics. by glucocorticoids (2.2–192 ng L−1), estrogens (5.6–60.2 ng L−1), This is similar with the situation of the studies in other countries and progestagens (6.9–40.5 ng L−1)(Liu et al., 2012b). (Salgado et al., 2010; Weigel et al., 2004). Pharmaceutical occurrence In Wuhan located in central China, study about seasonal variations in STPs of China is similar with the cases from Korea (Behera et al., of hormone contents in sewage treatment plants indicated contents 2011), Japan (Nakada et al., 2006), Finland (Vieno et al., 2006), and of natural E1 and E3 tended to be higher in winter than in summer, U.S. (Thomas and Foster, 2005), while higher contents up to probably due to their slow decomposition rate at low temperature. 482,687, 106,490, and 126,000 ng L−1 were found in the STP sewage The natural (E1, E2, and E3) and synthetic (diethylstilbestrol (DES)) samples of U.K. (Kasprzyk-Hordern et al., 2009), Portugal (Salgado et estrogens were believed to be the main contributors of the total estra- al., 2010), and Norway (Weigel et al., 2004), respectively. diol equivalents (EEQs), the index of estrogenic activity of sewage sample which is calculated directly by comparing the EC50 (concentration for 50% of maximal effect) of E2 and the sewage sample (Jin 3.1.1.4. Personal care products. Personal care product groups, e.g. anti- et al., 2008). However, it has been noted that other environmental microbial agents, synthetic musks, insect repellents, and UV filters, hormones, such as bisphenol A and alkylphenols, mainly caused the existed in STP sewage with concentration range of ND-3430 ng L−1 estrogenic effect of sewage (Sun et al., 2008), indicating the main for most reviewed cases in China. Triclosan and triclocarban were contributor of sewage estrogenic effect is probably case-specific. the most frequently detected antimicrobial agents (Zhao et al., The study in Beijing indicated estrogen concentrations in different 2010a), and HHCB and AHTN were the most frequently detected syn- effluents were in the order of pharmaceutical factory and STP thetic musks (Zeng et al., 2007). Extremely high contents of polycy- inlets N hospital N hennery N chemical factory N fish pool, suggesting clic musks up to 595.48 μgL−1 could be found in the cosmetic plant the main sources of hormones may rely on municipal wastewater and sewage (Chen et al., 2007). Input quantities of HHCB and AHTN into pharmaceutical-producing factory wastewater (Zhou et al., 2011e). STPs were estimated to be 0.2 and 0.06 g y−1 per capita of Shanghai, Investigation about androgens, progestogens, and estrogens in sew- and their environmental loads into receiving water were estimated to age of STPs confirmed androgens dominated in both influent and ef- be 1.26 and 0.38 t in 2007 (Zhang et al., 2008). The widely used insect fluent samples, followed by progestogens and estrogens (Chang et repellent DEET together with synthetic musks existed in Beijing STPs al., 2011). (Sui et al., 2011), and the sunscreen UV filters were present in Tianjin Levels of hormones in sewage samples of China are comparable STPs in North China (Li et al., 2007). with other areas, and are generally below the μgL−1 grade (Behera Wastewater levels of studied PCP groups are up to the μgL−1 et al., 2011; Nakada et al., 2006; Salgado et al., 2010; Servos et al., grade for most cases in China, while extremely high contents can be 2005). Relatively more steroid estrogens with contents exceeding found in manufacturing factories. Some groups of PCPs have received 1000 ng L−1 were found in several studies, such as Beijing in China more attention, e.g. polycyclic musks and triclosan, for their wide ap- (ND-4100 ng L−1)(Zhou et al., 2011e) and five cities in Portugal plication in the personal care products. The PCP levels in STP sewage (ND-2484 ng L−1)(Salgado et al., 2010). Both natural and synthetic of several other countries, such as U.K. (Kasprzyk-Hordern et al., hormones were detectable in sewage of China, indicating both natural 2009), Portugal (Salgado et al., 2010), and U.S. (Reiner et al., release and therapy use can contribute to the presence of hormones 2007a), could reach more than 10,000 ng L−1, exceeding the levels 214 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 in China, which may be attributed to the low per capita consumption than 20%, for carbamazepine during active sludge treatment (Miège of personal care products by the Chinese (Wang et al., 2012). et al., 2009). As to PCPs, disinfectants (triclocarbon, triclosan, and 2-phenylphenol) and preservatives (parabens) were greatly removed 3.1.2. Fates of PPCPs in STPs (Yu et al., 2011b), while polycyclic musks were partially removed by Mainly two processes are responsible for the fates of PPCPs in 26% to 70% with adsorption onto sludge as the dominant mechanism STPs, including biodegradation and adsorption by active sludge. (Zhou et al., 2009a). Biodegradation performance of PPCPs by active sludge can be Besides the species-specific removal performance of PPCPs, differ- affected by several factors, such as initial substrate concentration, ent STPs had varied removal for individual compounds owe to their temperature, and biodegradation process type. Biodegradation of different operation conditions (Zhou et al., 2010). Different treatment four antibiotics decreased when their initial influent concentrations processes also perform differently. Sui et al. (2010) found the target increased, probably due to the inhibition effect of microbial activities compounds, pharmaceuticals and DEET, could not be eliminated by by antibiotics (Yu et al., 2009). Biodegradation of synthetic musks sand filtration, but could be effectively removed by the advanced pro- showed better performance in summer than in fall which may be cesses—ozonation and microfiltration/reverse osmosis (MF/RO). attributed to the effect of temperature (Lv et al., 2010). Aerobic and From the studies above, it can be concluded some groups of PPCPs, anaerobic biodegradation have varied influence on different types of such as fluoroquinolones, hormones, caffeine, antimicrobial agents, PPCPs, such as diclofenac diminished mainly through anaerobic bio- and preservatives, can be removed efficiently with biodegradation degradation, while aerobic biodegradation governed the removal of and adsorption as the main mechanisms. However, some other anti-inflammatory drugs (ibuprofen, naproxen, indomethacin) and groups, such as majority of macrolides and sulfonamides, penicillin, lipid regulators (bezafibrate, clofibric acid, and gemfibrozil) (Huang fluconazole, and carbamazepine, are refractory to the conventional et al., 2011). Performance comparison of three kinds of biological sewage treatment processes and subsequently discharged unchanged treatment processes, including conventional active sludge (CAS), to the receiving water bodies. For the treatment of these PPCPs, biological nutrient removal (BNR), and membrane bioreactor (MBR), advanced technologies, such as oxidation (ozonation) and filtration revealed MBR resulted in better performance for the biodegradable (microfiltration and ultrafiltration) methods, may need to be introduced. species, e.g. caffeine and bezafibrate (Sui et al., 2011). Active sludge can also act as adsorbent for PPCPs in STPs. Majority of 3.1.4. PPCPs in sludge the antimicrobial agents (66–82% for triclosan and 82–86% for Through adsorption, groups of PPCPs can be found in sludge sam- triclocarban) were absorbed onto sludge (Yu et al., 2011b). Partitioning ples of STPs in China, including antibiotics (ND-21,000 ng g−1 dry may act as the dominant mechanism for adsorption, such as for the case weight (dw)) (Gao et al., 2012; Xu et al., 2007a), azole antifungal of 17β-estradiol (Yu et al., 2011b), while lipophilicity governed the drugs (ND-1442 ng g−1 (dw)) (Huang et al., 2010, 2012), hormones adsorption of antifungal drugs, e.g. clotrimazole, econazole, and micon- (1.6–372 ng g−1 (dw)) (Liu et al., 2011b), other pharmaceuticals azole (Peng et al., 2012). Adsorption process can be affected by pH, such (1.7–33.7 ng g−1 (dw)) (Yu et al., 2011a), antimicrobial agents as for the case of sulfonamide antibiotics (Yang et al., 2011). Meanwhile, (200.1–5088.2 ng g−1 (dw)) (Yu et al., 2011a), polycyclic musks desorption was observed reversibly from the sludge, implying sorbed (700–17,000 ng g−1 (dw)) (Zhou et al., 2009a), and sunscreen UV fil- PPCPs can be introduced into the environment if no further treatment ters (ND-24,700 ng g−1 (dw)) (Ruan et al., 2012), with concentra- was employed to remove them from the biomass (Yang et al., 2011). tions generally between ng g−1 and μgg−1 (dw). These studies Besides the two processes above, it was also found disinfection reveal PPCPs can be removed by adsorption onto sludge without process could eliminate some kinds of antibiotics, such as ampicillin being completely metabolized or degraded. This also demonstrates and cefalexin (Li and Zhang, 2011). the need to describe the composition and fate of PPCPs in sludge, since about half of the sludge is land applied and thus becomes a po- 3.1.3. Removal efficiencies of PPCPs in STPs tentially ubiquitous nonpoint source of PPCPs into the environment PPCPs can be removed in STPs mainly through the two processes, (Kinney et al., 2006). biodegradation and adsorption. Studies about PPCP removal in STPs To summarize the studies on PPCPs in STPs, groups of PPCPs can be suggest different groups of PPCPs perform differently. Antibiotics found in STP sewage as well as in sludge samples, including the had wide removal range up to 81%. Fluoroquinolones could be partial- normally detected antibiotics and hormones as well as other pharmaceu- ly removed with adsorption onto active sludge and suspended partic- tical species (such as caffeine, ibuprofen, and carbamazepine) and per- ulates as the possible mechanism, while majority of macrolides, sonal care products (such as triclosan, HHCB, and AHTN) in regions sulfonamides, penicillin, and imidazole could not be affected by the such as Pearl River Delta, Yangtze River Delta, and Beijing area. In China, treatment (Lin et al., 2009; Peng et al., 2011; Xu et al., 2007a). For large volume of wastewater containing PPCPs flows into sewage treat- azole antifungal drugs, fuconazole passed through the treatment ment plants (SIC, 2011). However, due to the limited removal efficiencies unchanged, ketoconazole could be readily biotransformed, while of some groups of PPCPs in STPs, they will be discharged into the receiving clotrimazole, econazole, and miconazole were more likely to be water and thus pose a potential risk to the aquatic environment. adsorbed onto the sludge (Peng et al., 2012). As a group of compounds with biodegradable properties, hormones can be eliminated efficiently. 3.2. PPCPs in surface water and sediments The most studied natural and synthetic steroid estrogens, including E1, E2, E3, and EE2, had high removal rates ranging from 83.2% to 98.8% After discharging from STPs, PPCPs in sewage would result in sub- (Zhou et al., 2011d), and the more abundant androgens and progesto- sequent contamination to the receiving aquatic environment. Groups gens also showed removal efficiency of 91–100% (Chang et al., 2011). of PPCPs have been detected in surface water bodies and sediments Studies in other countries confirmed norfloxacin (the fluoroquinolone) from South to North China, including rivers, lakes, and sea, even in as well as 17β-estradiol and estriol (the steroid estrogens) had removal the raw water sources of drinking water treatment plants (Table 3). rates of N80% in STPs (Miège et al., 2009). For other groups of pharmaceuticals, caffeine, the pharmaceutical with abundant occurrence in 3.2.1. PPCPs in surface water many cases, could be removed almost completely (Zhou et al., 2010), iopromide (one kind of X-ray contrast agents) could be greatly removed, 3.2.1.1. Antibiotics. Antibiotics occurred in surface water bodies with β-blockers were moderately eliminated, while carbamazepine, the concentrations generally below the μgL−1 grade (ND-776 ng L−1) antiepileptic drug, was not affected by the treatment (Yu et al., 2011b). for most reviewed cases in China. In Pearl River Delta, antibiotics Studies in other countries also found the low removal efficiency, less existed in Dongjiang River which serves as the drinking water source J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 215

Table 3 Occurrence of PPCPs in surface water.

Location Chemical Concentration Media Reference

Dongjiang River Antibiotics 0.9–67.4 ng L−1 Drinking water source Zhang et al. (2012a) Pearl River Antibiotics b35–510 ng L−1 River water Peng et al. (2008a) Pearl River Antifungal drugs b1–6.6 ng L−1 River water Huang et al. (2010) ⁎ −1 Pearl River Antibiotics ND−636 ng L River water Xu et al. (2007b) Victoria Harbor Antibiotics Below the limit of Seawater Xu et al. (2007b) quantiﬁcation (LOQ) Sea near Hong Kong Antibiotics b2–486 ng L−1 Seawater Gulkowska et al. (2007) Beibu Gulf Antibiotics ND*−50.9 ng L−1 Seawater Zheng et al. (2012) Jiulongjiang Antibiotics 0.05–775.5 ng L−1 River water Zhang et al. (2012c) Huangpu River Antibiotics ND−313.4 ng L−1 River water Jiang et al. (2011) Qiantang River Antibiotics 7.0–51.6 ng L−1 River water Tong et al. (2011) Yellow River and its tributaries Antibiotics b1–327 ng L−1 River water Xu et al. (2009b) Rivers near Beijing Antibiotics 1.3–535 ng L−1 River water Xiao et al. (2008) Hai River Antibiotics 26–210 ng L−1 River water Luo et al. (2011) Bo Sea Bay Antibiotics ND−6800 ng L−1 Seawater Zou et al. (2011) Bo Sea and Yellow Sea Antibiotics ND−16.6 ng L−1 Offshore seawater Zhang et al. (2013) Laizhou Bay Antibiotics ND−330 ng L−1 Seawater Zhang et al. (2012b) Dongjiang River Hormones ND−3.6 ng L−1 Drinking water source Gong et al. (2009) Pearl River Hormones 2.5–8.2 ng L−1 River water Gong et al. (2009) Pearl River Hormones ND−65 ng L−1 River water Peng et al. (2008b) Pearl River Hormones ND−75 ng L−1 River water Zhao et al. (2009) Danshui River Hormones ND−55.3 ng L−1 River water Liu et al. (2011b) Yundang Lagoon Hormones 0.3–6.8 ng L−1 Lake water Zhang et al. (2010) 0.7–6.4 ng L−1 Pore water Licun River near Jiaozhou Bay Hormones ND−180 ng L−1 River water Zhou et al. (2011b) Wulo Creek Hormones ND−1267 ng L−1 River water Chen et al. (2010) (Taiwan) Pearl River Pharmaceuticals 11.2–102 ng L−1 River water Zhao et al. (2010b) Pearl River Pharmaceuticals ND−490 ng L−1 River water Zhao et al. (2009) Pearl River Pharmaceuticals ND−2717 ng L−1 River water Peng et al. (2008b) Yangtze River Carbamazepine ND−1090 ng L−1 River water Zhou et al. (2011c) Rivers near Beijing Pharmaceuticals b24–7051 ng L−1 River water Zhou et al. (2010) Yellow River Pharmaceuticals ND−416 ng L−1 River water Wang et al. (2010a) Hai River ND−127 ng L−1 Liao River ND−717 ng L−1 (North China) Dongjiang River Antimicrobial agents b0.7–269 ng L−1 Drinking water source Zhao et al. (2013) Pearl River Antimicrobial agents 1.5–478 ng L−1 River water Zhao et al. (2013) Pearl River Antimicrobial agents b3.9–478 ng L−1 River water Zhao et al. (2010a) Pearl River Triclosan 0.6–347 ng L−1 River water Zhao et al. (2009) Pearl River Antimicrobial agents #NQ−2506 ng L−1 River water Peng et al. (2008b) Preservatives NQ−3142 ng L−1 Suzhou Creek Polycyclic musks 8–93 ng L−1 River water Zhang et al. (2008) (Shanghai) Yellow River Antimicrobial agents ND−64.7 ng L−1 River water Zhao et al. (2013) Hai River ND−117 ng L−1 Liao River b0.8–404 ng L−1 (North China)

Other counties Vienam Antibiotics 7–360 ng L−1 River water Managaki et al. (2007) (Mekong Delta) France Antibiotics ND-544 ng L−1 River water Tamtam et al. (2008) (Seine River) U.K. Antibiotics b0.5–183 ng L−1 River water Kasprzyk-Hordern et al. (2009) (Taff and Ely River) Finland Antibiotics b1.6–36 ng L−1 Drinking water source Vieno et al. (2007a) (Vantaa River) U.S. Antibiotics ND-694 ng L−1 River water Arikan et al. (2008) (Choptank River) U.S. Antibiotics ND-300 ng L−1 River water Kolpin et al. (2004) (Streams in Iowa) U.S. Antibiotics ND-1900 ng L−1 River water Kolpin et al. (2002) (139 streams) Korea Hormones 1.7–5.0 ng L−1 River water Kim et al. (2007) (Youngsan River) Spain Hormones 2–5ngL−1 River water Brix et al. (2009) (Llobregat River) Netherlands Hormones 0.4–10 ng L−1 River water Noppe et al. (2007) (Scheldt estuary) U.S. Hormones ND-20 ng L−1 River water Arikan et al. (2008) (Choptank River) U.S. Hormones ND-872 ng L−1 River water Kolpin et al. (2002) (139 streams) Hormones 0.03–18.9 ng L−1 River water Ferguson et al. (2013)

(continued on next page) 216 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224

Table 3 (continued) Location Chemical Concentration Media Reference

Australia (Little River) Other counties Korea Pharmaceuticals 1.1–361 ng L−1 River water Kim et al. (2007) (Youngsan River) Japan Pharmaceuticals ND–749 ng L−1 River water Nakada et al. (2008) (37 rivers and Tamagawa estuarine) U.K. Pharmaceuticals ND–5970 ng L−1 River water Kasprzyk-Hordern et al. (2009) (Taff and Ely River) U.K. Pharmaceuticals b1–928 ng L−1 River water Thomas and Hilton (2004) (5 rivers) Finland Pharmaceuticals 3–107 ng L−1 Drinking water source Vieno et al. (2007a) (Vantaa River) U.S. Pharmaceuticals ND–1950 ng L−1 River water Kolpin et al. (2004) (Streams in Iowa) U.S. Pharmaceuticals ND–10000 ng L−1 River water Kolpin et al. (2002) (139 streams) Brazil Pharmaceuticals 20–500 ng L−1 River water Stumpf et al. (1999) (Rivers in Rio de Janeiro State) India Triclosan Max. 5160 ng L−1 River water Ramaswamy et al. (2011) (Kaveri, Vellar, and Tamiraparani River) U.K. Antimicrobial agents b1–358 ng L−1 River water Kasprzyk-Hordern et al. (2009) (Taff and Ely River) Preservatives b0.2–305 ng L−1 UV ﬁlters b0.3–323 ng L−1 U.S. Synthetic musks 0.03–4.7 ng L−1 Lake water Peck and Hornbuckle (2004) (Michigan Lake) U.S. HHCB and ANTH ND–1200 ng L−1 River water Kolpin et al. (2004) (Streams in Iowa) Triclosan ND–140 ng L−1 DEET ND–130 ng L−1 ⁎ ND: not detected. #NQ: detected but. too low to be quantiﬁed.

− − for Hong Kong and Guangdong Province in southern China, with con- below the μgL 1 grade (ND-694 ng L 1). In China antibiotics up to centration range of 0.9–67.4 ng L−1. Macrolides in Dongjiang River 6800 ng L−1 were found in seawater of Bo Sea Bay (Zou et al., 2011). were probably related to domestic sewage discharge, while sulfon- Kolpin et al. (2002) measured 95 kinds of organic wastewater contami- amides and trimethoprim might be attributed to livestock agriculture nants in water samples from a network of 139 streams across 30 states wastewater (Zhang et al., 2012a). Xu et al. (2007b) compared the of U.S. during 1999 and 2000, and antibiotic contents in these streams − antibiotic levels in Pearl River of Guangzhou and its nearby Victoria could reach 1900 ng L 1. Harbor of Hong Kong in Pearl River Delta. The antibiotic concentrations in the coastal water of Victoria Harbor were generally below 3.2.1.2. Hormones. Hormones occurred in surface water of China with the limit of quantification (LOQ) while target analytes were all detect- content level generally below μgL−1 (ND-180 ng L−1)formost able in Pearl River, implying possible more serious contamination of reviewed studies, and also relatively lower than the antibiotic group. Pearl River by antibiotics. Antibiotic presence in the vicinity of aqua- They were also found in the drinking water source—Dongjiang River culture in Beibu Gulf suggested the contribution of aquaculture activ- with concentration up to 3.6 ng L−1 (Gong et al., 2009). In two reservoirs ities to antibiotic contamination in coastal water (Zheng et al., 2012). of Shenzhen, the total estrogen levels were in the range of 3–11 ng L−1 For Huangpu River in Yangtze River Delta, more veterinary antibi- (Lee et al., 2006). The Wulo Creek in Taiwan had hormone occurrence up otics, such as tetracyclines, were monitored in the suburban sampling to 1267 ng L−1 due to livestock feedlot discharge (Chen et al., 2010). sites compared with the urban sampling sites, confirmingtheroleoflive- In Beijing, hormone groups, e.g. estrogens, androgens, progestogens, stock wastewater as antibiotic contamination source (Jiang et al., 2011). glucocorticoids, and mineralocorticoids, were analyzed in 45 urban riv- The Hai River in North China contained veterinary antibiotics owe to ers, and it was found similar with the situation of sewage, androgens the release of swine farms and fishponds (Luo et al., 2011). In the nearby dominated among these groups, followed by glucocorticoids, progesto- coastal water of Bo Sea Bay, high density of human activities in the north gens, and estrogens. The principal component analysis for investigating bay resulted in more serious antibiotic contamination than the south the main contributor of urban river water contamination suggested bay (Zou et al., 2011). Antibiotics also occurred in the offshore water 62.7% of the mean summed hormones resulted from untreated sewage, of Bo Sea and Yellow Sea with concentration of ND-16.6 ng L−1,and 29.4% from treated sewage and/or naturally attenuated untreated sew- the risk assessment based on risk quotients (RQs) showed medium to age, and 7.9% from an unknown source, possibly pharmaceutical low ecological toxicity (0.01 b RQ b 1) derived from sulfamethoxazole, manufacturing plants (Chang et al., 2009). dehydration erythromycin, and clarithromycin to some sensitive aquatic These studies reveal the occurrence of hormones in surface water organisms, including Synechococcus leopoliensis and Pseudokirchneriella bodies as well as drinking water sources by the effect of wastewater dis- subcapitata (Zhang et al., 2013). charge, while untreated wastewater is probably the main contributor in These studies confirm that wastewater discharge is the main source of some cases. For other countries, rivers in most reviewed cases had reg- surface water pollution by antibiotics. Similar with the situation of sew- istered less than 20 ng L−1 of hormones (Arikan et al., 2008; Brix et al., age, contribution from livestock agriculture wastewater discharge is im- 2009; Ferguson et al., 2013; Kim et al., 2007; Noppe et al., 2007), except portant for antibiotic contamination. Surface water in other countries for the nationwide investigation in U.S. (ND-872 ng L−1)(Kolpin et al., seemed to be contaminated to a similar extent compared with the situa- 2002). tion of China, such as Vienam (Managaki et al., 2007), France (Tamtam et al., 2008), U.K. (Kasprzyk-Hordern et al., 2009), Finland (Vieno et al., 3.2.1.3. Other pharmaceuticals. Pharmaceutical groups, including anti- 2007a), and U.S. (Arikan et al., 2008; Kolpin et al., 2004), generally inflammatory drugs (e.g. salicylic acid, ibuprofen, diclofenac, J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 217 mefenamic acid, naproxen), blood lipid regulators (e.g. clofibric about the environmental fates of different dissociation species is acid, gemfibrozil), and antiepileptic drugs (e.g. carbamazepine), therefore suggested (Wei et al., 2013). showed widespread occurrence in the rivers from South to North China, such as Pearl River (Peng et al., 2008b), Yangtze River 3.2.3. PPCPs in sediments (Zhou et al., 2011c), Yellow River, Hai River, and Liao River (Wang PPCPs can be found in sediments of surface water through adsorp- et al., 2010a), with concentration range of ND-7051 ng L−1. Several −1 tion, but with much lower levels compared with sludge samples. Antibi- cases found pharmaceuticals more than 1000 ng L in the water − otics (ND-1560 ng g 1 (dw)) (Yang et al., 2010), antifungal drugs bodies (Peng et al., 2008b; Zhou et al., 2010, 2011c). Similar with (1–35 ng g−1 (dw)) (Huang et al., 2010), hormones (b0.9– the situation of sewage, caffeine dominated in some cases (Zhou − 10.9 ng g 1 (dw)) (Gong et al., 2011), antimicrobial agents (b0.6– et al., 2010). Detection frequency and median concentration of 2633 ng g−1 (dw)) (Yu et al., 2011a; Zhao et al., 2010a), polycyclic pharmaceuticals appeared higher during the low-flow season than − musks (b0.3–121 ng g 1 (dw)) (Zeng et al., 2008) existed in the the high-flow season, probably due to the dilution effect by rainfall sediments of Pearl River and coastal water in southern China, antibi- (Peng et al., 2008b). Sites near metropolitan areas appeared more − otics (ND-5770 ng g 1 (dw)) occurred in the sediments of Yellow contaminated (Wang et al., 2010a), and the pharmaceutical pollu- River, Hai River, and Liao River in northern China (Zhou et al., tion may be closely associated with the direct discharge of − 2011a), and UV filters (0.3–7.1 ng g 1 (dw)) were present in sedi- untreated wastewater (Yu et al., 2011b). The contamination level ments of Songhua River in northern China (Zhang et al., 2011c), of surface water by pharmaceuticals in other countries seemed suggesting the PPCP adsorption on sediments as well as the possible comparable with China, such as Korea (Kim et al., 2007), Japan desorption. For better understanding PPCPs in sediments as a poten- (Nakada et al., 2008), U.K. (Kasprzyk-Hordern et al., 2009; Thomas tial pollution source to surface water, more studies about PPCPs in and Hilton, 2004), U.S. (Kolpin et al., 2002; Kolpin et al., 2004), sediments are needed. and Brazil (Stumpf et al., 1999), with concentration range of ND-10,000 ng L−1. Pharmaceuticals up to 107 ng L−1 were found in Vantaa River as drinking water source in Finland (Vieno et al., 3.2.4. Removal efficiencies of PPCPs in water treatment plants 2007a). About the removal efficiency of PPCPs in water treatment plants, Qiao et al. (2011) stated conventional water treatment processes 3.2.1.4. Personal care products. Personal care products in surface water (coagulation and flocculation) removed the type and concentration of China had an occurrence range of ND-3142 ng L−1, while contents of PPCPs by 30% and 50% respectively, while advanced treatment of most cases were below the μgL−1 level. Urban rivers of Guangzhou processes could achieve the reduction ratios of 50% and 90% respec- in Pearl River Delta contained disinfectants (triclosan, triclocarban, tively. Study in Korea confirmed that conventional treatment had 2-phenylphenol) and preservatives (methyparaben, propylparaben) limited PPCP removal, however, granular activated carbon adsorp- with high detection frequency and concentration (Peng et al., tion could result in efficient control (Kim et al., 2007). Another 2008b; Zhao et al., 2009). Triclosan and triclocarban in rivers, includ- study also found the conventional coagulation and sedimentation ing Dongjiang River, Pearl River, Yellow River, Hai River, and Liao showed much marginal removal of studied pharmaceuticals by 3%, River, had high occurrence in sites located near the urban area and sand filtration could achieve more additional removal of 10%, while were significantly influenced by the factors such as the total or the advanced oxidation method, ozonation, reduced most pharma- untreated urban domestic sewage discharge at river basin scale ceuticals from up to 107 ng L−1 to below the detection limits of (Zhao et al., 2013). Polycyclic musks occurred in Suzhou Creek of 0.2–7.0 ng L−1 (Vieno et al., 2007a). Shanghai in Yangtze River Delta with HHCB and AHTN as the two To summarize the studies about PPCPs in surface water bodies, main species (Zhang et al., 2008). When looking at the PCP studies different groups of PPCPs occur ubiquitously in aquatic environment worldwide, Pearl River in China with antimicrobial agents up to from South to North China, owing to the impact of municipal and ag- 2506 ng L−1 and preservatives up to 3142 ng L−1 (Peng et al., riculture wastewater discharge. Antibiotics and steroid estrogens can 2008b) and Kaveri, Vellar, and Tamiraparani River in India with triclo- even be found in source water of drinking water treatment plants. san maximum concentration of 5160 ng L−1 (Ramaswamy et al., The polluted surface water can be attributed to the huge wastewater 2011) appeared to be more contaminated by PCPs. discharge volume and limited wastewater treatment rate in current China. The discharge volume of municipal, industry, and agriculture 3.2.2. Fates of PPCPs in surface water wastewater is huge as well as with a rapid growth rate. In 2009, In natural surface water bodies, PPCPs may undergo several 59 billion tonnes of wastewater were discharged into the receiving environmental processes, mainly adsorption by sediments and pho- aquatic environment of China. From 2003 to 2009, the municipal sew- tolysis. Majority of antimicrobial agents, such as triclosan, appeared age discharge volume increased by 40% (SIC, 2011). However, the to be adsorbed onto sediments. Pores of sediments can capture tri- wastewater treatment rate is relatively low due to the lack of facili- closan, while the sediment organic matter may bound triclosan and ties, compared with the developed countries. Wastewater treatment reduce its solubility. The adsorption process can be affected by pH rate in rural regions of China was as low as 60% in 2010 (MOHURD, (Lin et al., 2011). Adsorption can remove PPCPs from the water bod- 2012), meaning that up to 40% of the wastewater was discharged ies, however, sediments may also become a sink for PPCPs which directly into the receiving water bodies without any treatment. The may be released back to the aquatic environment (Zhao et al., current wastewater discharge and treatment condition finally results 2013). in the pollution of surface water by PPCPs in China. Water pollution is Aqueous photolysis was observed for tetracycline, and the pro- a large challenge for the developing China, and numerous efforts need cess could be enhanced by high pH (Jiao et al., 2008). The non- to be made to solve this problem in the future. steroidal anti-inflammatory drug diclofenac can also be degraded Removal efficiency of PPCPs in water treatment plants is not satis- by sunlight. pE value of the aqueous environment may affect the factory by the conventional treatment processes, and thus advanced forms of N and further influent the photolysis process (Zhang et al., steps, e.g. ozonation and activated carbon adsorption, need to be in- 2011a). Further photolytic mechanism study for the different disso- troduced for the control of PPCP problem in source water. Besides ciation species of ciprofloxacin indicated five dissociation species surface water as drinking water source, the studies about PPCPs in had dissimilar photolytic kinetics, including stepwise cleavage of groundwater are urgent. Meanwhile, more studies about PPCPs in thepiperazinering,defluorination, and oxidation. For accurate seawater which possesses varied physical, chemical, and biological ecological risk assessment of ionizable emerging pollutants, study properties with river water are needed. 218 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224

3.3. PPCPs in soil 4. Human exposure and body loading of PPCPs

PPCPs can be introduced into soil through sludge land applica- 4.1. Human exposure tion or landfill, use of livestock wastes as fertilizers, and reclaimed water irrigation. Pollutants in soil may be accumulated in plants or Humans can be exposed to PPCPs mainly through daily use of migrate through soil intact or transformed and reach groundwater, personal care products, inhalation, and dietary intake. More than finally resulting in pollution to the drinking water source. Veteri- 80% of the investigated daily personal care products, including nary antibiotics existed in the soil from organic vegetable farmland tooth pastes, hair care products, body washes, soaps, skin lotions, fertilized with livestock wastes in Tianjin (Hu et al., 2010a), with and makeup, contained synthetic musks with total contents as concentrations up to 2683 ng g−1 (dw). Azole biocides were high as 1.02 mg g−1. HHCB was the predominant species in all of detected in biosolid-amended soil and showed persistence with the personal care products analyzed, accounting for 52% of the dissipation half lives up to 440 days (Chen et al., 2013). Chen et total musk concentrations on average. The dermal exposure rates al. (2011) found irrigation with wastewater could result in more in adults were calculated to be 3.38 mg d−1 for musks (Lu et al., PPCPs in soil compared irrigation with groundwater. The study 2011). about sorption and degradation behavior of drugs in soil discov- For the low volatility of PPCPs, their participation into atmospheric ered some drug species, such as diclofenac and ibuprofen, showed environment would be limited. However, some PPCPs, mainly groups poor adsorption as well as with persistence under anaerobic condi- with relatively high volatility, were still found in indoor dusts or air. tion (Lin and Gan, 2011). The simulated turf grass irrigation for the Siloxanes which are widely used in personal care and other consumer study of PPCP leachability discovered estrone, ibuprofen, naproxen, products were found in the indoor dust samples with total concentra- clofibric acid, and triclosan in the 30-cm soil profiles, indicating tion range of 21.5 to 21,000 ng g−1. Their concentrations in indoor the potential to contaminate groundwater. The leachability of dusts were associated with the number of electrical/electronic appli- PPCPs may be affected by the chemicals' characteristics (such as ances and number of occupants and smokers living in the house. Tod- pKa values), properties of soil (such as soil organic matter and dlers had higher estimated daily intake than adults (Lu et al., 2010b). clay content), and salinity of the irrigation water (Xu et al., Another study investigated the occurrence of two polycyclic musks, 2009a; Xu et al., 2010). These studies reveal PPCPs can be intro- three nitro musks, and one metabolite of HHCB (HHCB lactone) in in- duced into soil by these routes. Some groups of PPCPs may per- door dust samples. The total concentrations of musks ranged from 4.4 form poor adsorption in soil and be refractory to biodegradation, to 688 ng g−1 with HHCB as the predominant species (Lu et al., suggesting their possible leaching risk. Therefore, reclaimed 2010a). In the polycyclic musk investigation from a typical cosmetic water irrigation and sludge and livestock waste land application plant in Guangzhou, except for sewage and sludge, polycyclic musks need to be evaluated carefully for the potential risk of PPCP were found to be mostly contained in the gas phase at the percentage introduction. of 86.4–97.7% (Chen et al., 2007). Parabens in indoor dusts from four countries, including U.S., China, Japan, and Korea, were investigated, 3.4. PPCPs in wild animals and paraben species with contents up to 110,800 ng g−1 were detected. Geometric mean concentrations of the investigated parabens The studies about PPCPs in wild animals of China are limited. in indoor dusts from four countries were in the order of Korea Fluoroquinolones were found in fish species from marine aquacul- (2320 ng g−1) N Japan (2300 ng g−1) N U.S. (1390 ng g−1) N China ture regions of the Pearl River Delta, including rabbitfish, black (418 ng g−1). The relatively lower parabens in dusts from China is porgy, and gray snapper, with liver concentration up to probably related to the low per capita consumption of personal care 254.6 ng g−1 wetweight(ww)(He et al., 2012). Steroid estrogens products by the Chinese (Wang et al., 2012). had contents up to 11.3 ng g−1 (dw) in wild fish species from PPCPs existed in drinking water and food as the dietary sources for Dianchi Lake in Southwest China, e.g. crucian carp, carp, and sil- humans. Four kinds of PPCPs existed in the drinking water from dis- very minnow. Liver had the highest estrogen accumulation, fol- tribution network with contents around 1 ng L−1 in South China lowing by gill and muscle (Liu et al., 2011a). Wan et al. (2007) (Qiao et al., 2011). Fluoroquinolone antibiotics also occurred in the investigated the distribution of musk fragrances in different or- tap water of Guangzhou and Macau (Wang et al., 2010b). Triclosan gans of Chinese sturgeons and found higher concentration of fra- was present in both tap water and bottled water with concentrations grances in adipose tissues (1.0–62.1 ng g−1 (ww)), indicating up to 14.5 and 9.7 ng L−1 respectively. Triclosan may also be released their distribution was associated with the affinity to lipid. The from baby bottles. Daily intakes of triclosan by adults and infants lipid-corrected concentrations increased when the age of Chinese were estimated to be 10 and 5 ng d−1 respectively (Li et al., 2010a). sturgeons increased, which is consistent with the situation of Vegetables contained antibiotics in the organic farm bases with live- fish, but different from that of mammals. Hu et al. (2011) discov- stock wastes as fertilizers in Tianjin of North China, with concentra- ered the biota-sediment accumulation factors (BSAFs) higher than tion up to 532 ng g−1. The antibiotics accumulated in vegetables 1.7 for two common polycyclic musks, HHCB and AHTN, in crucian probably through water transport and passive absorption, and the carp, common carp, and silver carp, revealing the possible distribution in the plant was in the order of leaf N stem N root (Hu bioaccumulation of these fragrances. The contents of synthetic et al., 2010a). musks in wild fish from China are comparable with that from Japan (b1.9–149 ng g−1 (ww)) (Nakata, 2005) and Denmark 4.2. Body loading (ND-52.6 ng g−1 (ww)) (Duedahl-Olesen et al., 2005). The study in Denmark also revealed the decreasing tendency of synthetic Through the different exposure pathways, PPCPs were found in musks in wild fish according to time from 1992 to 2004 kinds of human samples, e.g. human milk, blood, and urine. Synthetic (Duedahl-Olesen et al., 2005). These studies suggest the presence musks existed in the human milk samples from the mothers living in and possible bioaccumulation of PPCPs in wild animals. However, Chengdu of Southwest China, with concentration of b1.4–16.5 ng g−1 the limited studies about PPCPs in wildlife of China so far are lipid weight. Their occurrence was related with the frequent use of per- mainly focused on antibiotics, hormones, and synthetic musks. sonal care products, such as hand-cleaning agents, body lotions, sham- Occurrence and behavior of other PPCP groups in wild animals poos, hair dyes, and hair gels (Yin et al., 2012). Four synthetic musks, need to be investigated in the future, and the potential adverse MX, MK, HHCB, and AHTN, were also present in the human milk sam- effects need to be evaluated. ples from three cities of eastern China with contents of 4–63 ng g−1 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 219 lipid weight. The dominant species was HHCB, followed by MX. Daily gene transfer (HGT) is a major mechanism for sharing ARGs between intake of synthetic musks by infants from milk was estimated to be microorganisms, and antibiotic resistance determinants can be trans- 277–7391 ng d−1 (Zhang et al., 2011b). The occurrence of synthetic mitted to other animals and humans via consumption of polluted musks in human milk of China seems relatively lower compared with water as well as agricultural and aquacultural products (Kobayashi the studies of U.S. (b2–917 ng g−1 lipid weight) (Reiner et al., et al., 2007). 2007b), Denmark (38–422 ng g−1 lipid weight) (Duedahl-Olesen et The situation with respect to antibiotic resistance derived from al., 2005), and Sweden (b2–268 ng g−1 lipid weight) (Lignell et al., antibiotic abuse in China is severe (Reynolds and McKee, 2009). In 2008). Two nitro musk fragrances and five polycyclic musk fragrances 2010, the national Ministry of Health collected more than 270,000 were measured in human blood samples from 11 cities of China. isolated bacteria samples from 128 hospitals across the country to HHCB had the highest content of 0.85 ng g−1 and detection frequency study the antibiotic resistance situation. Results showed the resis- of 91%, followed by AHTN with the highest content of 0.53 ng g−1 tance rates had reached nearly 80%, creating a severe situation for and detection frequency of 77% (Hu et al., 2010b). The study on syn- public health (Chinadaily, 2012). The level of antibiotic resistance in thetic musks in human blood plasma of young adults from Austria dis- China was compared with Kuwait and U.S., revealing China had the covered the dominant HHCB and MX with median concentrations of highest value. China also had the most rapid growth rate of resistance 420 and 11 ng g−1, respectively (Hutter et al., 2009), which are signif- (22% average growth in a study period of 1994 to 2000), while Kuwait icantly higher than the previous study in China. Triclosan was present in was second (17% average growth from 1999 to 2003), and U.S. was the urine samples of 287 children and students aged from 3 to 24 years the lowest (6% average growth from 1999 to 2002) (Zhang et al., old in Guangzhou, with detection rate of 93% and mean concentration of 2006). ARGs were widely found in the aquatic environment of 3.55 μgg−1 creatinine (3.77 μgg−1)(Li et al., 2013). The occurrence of China. Tetracycline resistance genes existed in sewage of wastewater parabens and the metabolites were investigated in the urines of U.S. and treatment plants in Hong Kong and Shanghai (Zhang et al., 2009a). Chinese people. Medium concentrations of the sum of 6 investigated Escherichia coli resistant to antibiotics, mainly sulfonamides, tetracy- parabens in urine of U.S. and Chinese children were 54.6 and cline, and ampicillin, existed in Wenyu River Basin in Beijing, while 10.1 ng mL−1 respectively, while the concentration for Chinese adults multiantibiotic resistance was also discovered (Hu et al., 2008). Sul- was 33.2 ng mL−1, which was relatively lower than the reported fonamide resistance genes occurred in Hai River. Their concentrations value for U.S. adults. Elevated urinary paraben concentration (1000– in sediments were 120–2000 times higher than that in water, indicat- 10,000 ng mL−1) were found for Chinese adult females (Wang et al., ing sediments are probably an important reservoir for ARGs in surface 2013). water (Luo et al., 2010). It is believed PPCPs pose no significant risk to human health based Hormones may result in the adverse effects of endocrine disrup- on current environmental exposure and human body burden data tion. These endocrine disruptors can have a great range of reproduc- (Touraud et al., 2011). However, the studies above demonstrate the tive and development effects, e.g. reduced fertility, feminization of accumulation of PPCPs in human bodies, and their occurrence in males, induction of vitellogenesis in male (plasma vitellogenin induc- human milk implies the possible impact to infants. Studies about tion), and intersex phenomena (Lai et al., 2002). These hormones PPCP human exposure and body loads are mainly about synthetic possess 10,000–100,000 times higher estrogenic potency compared musks, antimicrobial agents, and parabens. Investigation about with exogenous endocrine disrupting chemicals (Khanal et al., other PPCP groups is urgent. Meanwhile, for the full understanding 2006), and are believed to be the most significant estrogenic com- of potential risks derived from PPCP exposure, a comprehensive in- pounds present in many sewage effluents (Jobling et al., 2006). A vestigation about the contribution of different exposure routes is number of aquatic species, for example, crucian carp, trout, minnow, needed. and turtle, have been reported to be sexually inhibited or reversed by the presence of estrogens at environmental levels (Irwin et al., 2001; 5. Toxicity and risk assessments of PPCPs Liu et al., 2012a; Tabata et al., 2001; Zha et al., 2008). For the personal care product group, parabens and UV filters can also act as endocrine 5.1. Toxicity disruptors (Gomez et al., 2005), while triclosan is suspected to result in the endocrine disrupting effects (Foran et al., 2000). The wide occurrence of PPCPs in different environmental media Table 4 lists some examples of adverse effects from different kinds raises concerns about their potential harm to ecosystem and human of PPCPs. Besides the antibiotic resistance by antibiotics and endo- health. Based on the toxicological data and environmental concentra- crine disruption by hormones, it was found antibiotics (Yang et al., tion levels, risks of acute toxic effects from these PPCPs are believed to 2008) and gemfibrozil (blood lipid regulator) (Rosal et al., 2009)as be not likely (Brausch and Rand, 2011; Fent et al., 2006). However, as well as triclosan and triclocarban (antimicrobial agents) (Yang et al., indicated by the case that residues of veterinary diclofenac caused the 2008) could inhibit the growth of algae. Caffeine (stimulant drug) significant decline of vulture population in Pakistan, severe adverse may result in endocrine disruption of goldfish (Li et al., 2012), and effects such as wildlife population decline may occur (Lindsay and propranolol (beta-blocker) can reduce the viable eggs of Japanese Martin, 2004). Meanwhile, chronic effects could not be excluded. medaka (Huggett et al., 2002). Carbamazepine (antiepileptic drug) The continuous input of PPCPs into the environment may accumulate and HHCB (polycyclic musk) may result in oxidation stress to rain- and cause irreversible harm to wildlife and human beings (Brausch bow trout (Li et al., 2010b) and goldfish (Chen et al., 2012). and Rand, 2011; Fent et al., 2006). Diclofenac, the nonsteroidal anti-inflammatory drug, may cause Environmental exposure to antibiotics may accelerate the persis- renal lesions and gill alterations to rainbow trout (Schwaiger et al., tence or emergence of antibiotic resistance genes (ARGs) which 2004). Furthermore, synergistic effects of these toxic chemicals can pose potential harm to ecosystem and human health (Kemper, be a concern. Tests with combinations of various pharmaceuticals 2008; Zhang et al., 2009b). Antibiotic resistance genes encoding resis- (carbamazepine, diclofenac, and ibuprofen) revealed stronger effects tance to a broad range of antibiotic species, such as macrolides, sul- than expected from the effects measured singly to the target aquatic fonamides, fluoroquinolones, and tetracyclines, occur ubiquitously organism—Daphnia magna (Cleuvers, 2003). Another concern about in hospital and livestock feeding effluents, municipal wastewater, PPCPs is their potential bioaccumulation and biomagnification in surface water, as well as drinking water resources (Pruden et al., aquatic food web. Several groups of PPCPs, such as UV filters, disinfec- 2006; Zhang et al., 2009b). On the other hand, multiple antibiotic tants, and synthetic musks, have been approved to be able to resistant (MAR) superintegrons, which may contain over 100 ARG bioaccumulate and probably biomagnify and may finally cause cassettes, have been also discovered (Pruden et al., 2006). Horizontal adverse effects to human beings (Brausch and Rand, 2011). 220 J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224

Table 4 Adverse effects of PPCPs.

Chemical Adverse effect Exposure dose Exposure span Reference

Roxithromycin, clarithromycin, tylosin Growth inhibition of algae (Pseudokirchneriella subcapitata)40–64 μgL−1 3d Yang et al. (2008) (Antibiotics) Caffeine Endocrine disruption of goldfish (Carassius auratus) 2000 μgL−1 7d Li et al. (2012) (Stimulant drug) Diclofenac Renal lesions and gill alterations of rainbow trout (Oncorhynchus mykiss)5μgL−1 28 d Schwaiger et al. (2004) (Nonsteroidal anti-inflammatory drug) Carbamazepine Oxidation stress of rainbow trout (Oncorhynchus mykiss) 200 μgL−1 42 d Li et al. (2010b) (Antiepileptic drug) Gemfibrozil Growth inhibition of algae (Anabaena sp.) 4420 μgL−1 1d Rosal et al. (2009) (Blood lipid regulator) Propranolol Reduction of viable eggs of Japanese medaka (Oryias latipes) 0.5 μgL−1 28 d Huggett et al. (2002) (β-blocker) HHCB Oxidation stress of goldfish (Carassius auratus) 1.5 μgL−1 7–21 d Chen et al. (2012) (Synthetic musk) Triclosan and triclocarban Growth inhibition of algae (Pseudokirchneriella subcapitata) 0.4–10 μgL−1 3d Yang et al. (2008) (Antimicrobial agents)

5.2. Risk assessments 6. Regulations and control strategies of PPCP contamination

Many studies about risk assessments of PPCPs base on the Europe- Regulations are established for PPCP management in several coun- an Medicines Agency (EMA) guidelines, in which the risk quotient is tries. In U.S. the assessment of environmental risks from pharmaceu- calculated as the ratio between predicted environmental concentra- ticals has been required by U.S. Food and Drug Administration tions (PEC) and predicted no-effect concentrations (PNEC). If risk (U.S.FDA) under the National Environmental Policy Act (NEPA) quotient is higher or equal to one, it suggests that the particular sub- since 1969 (U.S.FDA, 1969), and in 1998 the guidance for a tiered stances could cause potential adverse ecological effects (EMA, 2006). risk assessment method was published by Center for Drug Evaluation Stuer-Lauridsen et al. (2000) evaluated the risks of 25 most used and Research (CDER) of U.S.FDA (CDER, 1998). In the same year U.S. pharmaceuticals in aquatic environment of Denmark, and PEC/PNEC Environmental Protection Agency (U.S.EPA) promulgated revised reg- ratios of ibuprofen, acetylsalicylic acid, and paracetamol were found ulations for the pharmaceutical industry to control both effluent dis- to exceed one, and the ratios of estrogens approached one. The risk charges and air emissions (U.S.EPA, 1998). For European Union (EU), assessment for most used drugs in aquatic environment of England in 2006, the first Guideline for Environmental Risk Assessment of found the PEC/PNEC ratios of paracetamol, amoxycillin, oxytetracy- Human Medicines was published by the European Medicines Agency cline, and mefenamic acid exceeded one (Jones et al., 2002). Besse (EMA, 2006). Switzerland, a non-member of EU, requires only and Garric (2008) evaluated 120 pharmaceuticals and 30 metabolites ecotoxicity information for human pharmaceuticals and both in the aquatic environment of France according to their biological and ecotoxicity and potential environmental risks for veterinary drugs physicochemical properties as well as predicted environmental con- (AMZV, 2001). In Australia, pharmaceutical environmental risk as- centrations, and 40 pharmaceuticals and 14 metabolites were sug- sessment is required with new medicine registration by Australian gested to need prioritized attention. Prioritization and ranking Therapeutic Goods Administration (TGA, 2008). methods are needed for screening target compounds for their poten- In China, the environmental management regulations for hazard- tial risks. Cooper et al. (2008) developed a risk assessment database ous chemicals include the Environmental Management Method for for commonly used pharmaceuticals based on their environmental New Chemicals promulgated in 2010 (MEP, 2010) and Environmental behavior and ecological toxicity focusing on the marine and Management and Registration Method for Hazardous Chemicals esturarine environment, and anti-infectives were found to pose the which comes into force recently in March 2013 (MEP, 2012). The highest risk. A prioritization approach based on number of prescrip- environmental risk assessments of drugs and cosmetics are not in- tions and toxicity information as well as human metabolism and cluded in these regulations, however, their ingredients and interme- wastewater treatment removal information was performed on 200 diates during production are regulated under these rules. The most-prescribed drugs in U.S. in 2009, finding levothyroxine and human health risk assessments of new drugs are required under the montelukast sodium received the highest scores (Dong et al., 2013). Pharmaceutical Administration Law (CFDA, 2001), however, environ- A ranking system based on the pharmaceutical consumption, removal mental risk assessments of drugs and cosmetics are not regulated cur- performance in the wastewater treatment plants, and potential eco- rently. The livestock agriculture waste discharge volume is regulated logical effects was applied to the situation of China. Erythromycin, by the national Ministry of Environmental Protection (MEP) of China diclofenac acid, and ibuprofen had the high priority among 39 phar- (MEP, 2001). Wastewater discharge volume of pharmaceutical manu- maceuticals. Antibiotics contributed most to the pharmaceuticals facturers producing 16 kinds of drugs, e.g. caffeine, ibuprofen, and with priority. However, only 32% of the antibiotics were listed as sulfadiazine, is also regulated (MEP, 2008). In the future, regulations priority drugs, compared with the 71% and 100% for anti-inflammatory about environmental risk assessment as well as the assessment and antilipidemic drugs. Thus, the author stated antibiotics might be guidelines for PPCPs should be established in China. Source control overanxiously considered (Sui et al., 2012). is a direct and efficient strategy. Expired drugs should be collected In general, based on the current environmental contamination through the establishment of a drug management and reclamation levels of PPCPs, it is believed the acute toxicity caused by PPCPs scheme. Removal performance of PPCPs in STPs as well as in drinking has low possibility, however, chronic toxicity could not be excluded. water treatment plants needs to be improved by the adoption of ad- The high level and rapid growth of antibiotic resistance in China vanced treatment technologies. In addition, sewage sludge landfill, derived from antibiotic abuse need a special attention. Current risk livestock waste land use, and reclaimed water irrigation need more assessments of PPCPs are mainly based on the acute toxicity data, careful evaluation considering possible introduction of PPCPs into more accurate evaluation considering chronic toxicity exposed at soil and groundwater environment. For the specific situation of environmental concentration levels is urgent. Moreover, more risk China, the antibiotic abuse needs to be improved in the future for assessments for personal care product groups should be conducted. the control of current serious antibiotic resistance. More wastewater J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224 221 treatment facilities need to be established to increase the current Brix R, Postigo C, González S, Villagrasa M, Navarro A, Kuster M, et al. Analysis and occurrence of alkylphenolic compounds and estrogens in a European river basin and an sewage treatment rate and avoid direct sewage discharge into aquatic evaluation of their importance as priority pollutants. Anal Bioanal Chem 2009;396: environment, especially in the rural areas of China. Protection of 1301–9. drinking water sources is not sufficient in current China. Management Brown KD, Kulis J, Thomson B, Chapman TH, Mawhinney DB. Occurrence of antibiotics in hospital, residential, and dairy effluent, municipal wastewater, and the Rio and supervision of drinking water sources need to be strengthened. Any Grande in New Mexico. Sci Total Environ 2006;366:772–83. possible pollution source of PPCPs, e.g. wastewater treatment plants, Carballa M, Omil F, Lema JM, Llompart M, Garcıa-Jareś C, Rodrıgueź I, et al. Behavior of livestock agriculture, hospitals, and cosmetic plants, should not be pharmaceuticals, cosmetics and hormones in a sewage treatment plant. Water Res – located in the vicinity of drinking water sources, and any illegal waste- 2004;38:2918 26. CAST (China Association of Science and Technology). The threat of antibiotic abuse to water discharge to the source water should be strictly forbidden. public safety; 2008. Castiglioni S, Bagnati R, Fanelli R, Pomati F, Calamari D, Zuccato E. Removal of pharmaceuticals in sewage treatment plants in Italy. Environ Sci Technol 2006;40:357–63. 7. Conclusions CDER (Center for Drug Evaluation, Research). Guidance for industry-environmental assessment of human drug and biologics applications. U.S. Department of Health and Human PPCPs are present ubiquitously in different environmental media Services, Food and Drug Administration, U.S. Center for Drug Evaluation and Research (CDER), Center for Biologics Evaluation and Research (CBER); 1998 [July 1998, CMC 6, of China. The contamination levels of PPCPs in sewage and surface Revision 1. http://www.fda.gov/downloads/Drugs/GuidanceComplianceRegulatory −1 −1 water are in the range of ng L to μgL , while the levels in sludge, Information/Guidances/ucm070561.pdf. Accessed by May 2013]. sediments, and soil are in the range of ng g−1 to μgg−1 (dw). With CFDA (China Food, Drug Administration). Pharmaceutical Administration Law of the People's Republic of China. http://www.sda.gov.cn/WS01/CL1030/23396.html, huge wastewater discharge volume in China, the environmental 2001. (Accessed by May 2013). loads of PPCPs are tremendous. The surface water pollution by Chang H, Hu JY, Shao B. Occurrence of natural and synthetic glucocorticoids in sewage PPCPs may be attributed to the huge wastewater discharge volume treatment plants and receiving river waters. Environ Sci Technol 2007;41:3462–8. Chang H, Hu JY, Wang LZ, Shao B. Occurrence of sulfonamide antibiotics in sewage and limited wastewater treatment rate in current China. For the pro- treatment plants. Chinese Sci Bull 2008;53:514–20. tection of surface water, there are great needs to improve the waste- Chang H, Wan Y, Hu JY. Determination and source apportionment of five classes of water treatment rate to reduce the direct discharge. In some cases, steroid hormones in urban rivers. Environ Sci Technol 2009;43:7691–8. PPCPs can be detected in source water of drinking water treatment Chang H, Wan Y, Wu SM, Fan ZL, Hu JY. Occurrence of androgens and progestogens in wastewater treatment plants and receiving river waters: comparison to estrogens. plants or even tap water. It is urgent to protect the water sources Water Res 2011;45:732–40. from PPCP contamination. Chang XS, Meyer MT, Liu XY, Zhao Q, Chen H, Chen JA, et al. Determination of antibi- There is severe lack of information about other environmental otics in sewage from hospitals, nursery and slaughter house, wastewater treatment plant and source water in Chongqing region of Three Gorge Reservoir in media contamination status of PPCPs in China, such as surface and China. Environ Pollut 2010;158:1444–50. ground drinking water sources, wildlife, as well as the human body Chen DH, Zeng XY, Sheng YQ, Bi XH, Gui HY, Sheng GY, et al. The concentrations and dis- burden. Areas other than the several hotspots all over China need to tribution of polycyclic musks in a typical cosmetic plant. Chemosphere 2007;66: 252–8. be investigated about their PPCP pollution status. Meanwhile, most Chen F, Gao J, Zhou QX. 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Chen ZF, Ying GG, Ma YB, Lai HJ, Chen F, Pan CG. Occurrence and dissipation of three Acknowledgments azole biocides climbazole, clotrimazole and miconazole in biosolid-amended soils. Sci Total Environ 2013;452–453:377–83. Chinadaily. http://www.chinadaily.com.cn/china/2012-03/23/content_14893869.htm, Financial support from Seed Collaborative Research Fund from 2012. [Accessed by May 2013]. – the State Key Laboratory in Marine Pollution (SCRF0003) and Na- ChinaIRN (China Industry Research Net). 2012 2013 personal care product market development analysis; 2012. tional Natural Science Foundation of China/Research Grants Council Cleuvers M. Aquatic ecotoxicity of pharmaceuticals including the assessment of combi- (N_HKBU210/11) are gratefully acknowledged. nation effects. Toxicol Lett 2003;142:185–94. Cooper ER, Siewicki TC, Phillips K. Preliminary risk assessment database and risk ranking of pharmaceuticals in the environment. Sci Total Environ 2008;398:26–33. 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